36.1
Chapter 36
Impacts of mining
Jon Harding and Ian Boothroyd
INTRODUCTION
New Zealand has a long history of mineral extraction,
beginning with the discovery of gold in the Coromandel
Peninsula in 1852 and Central Otago in the 1861, and
coal on the West Coast in the 1860s. Coal was first mined
in 1872 at the Albion mine, and soon after underground
mines were extensively developed along the StocktonDenniston Plateau, North Westland, and in the Reefton
area, where coal was discovered in close proximity to gold.
In Central Otago widespread alluvial gold mining began
in the 1860s, while the Ohinemuri goldfield opened in
the Coromandel in 1875. Most of the early activity
occurred in the Karangahake area, but in 1878 the rich
gold-bearing quartz at Martha Hill in Waihi was
discovered. Placer mining and gold dredging operations
became widespread in Central Otago and Westland in the
1870s and these areas continue to be reworked as both
technology and the price of gold improve (Fig. 36.1).
Other minerals that have been mined in New Zealand
include gravel, tin, copper and uranium.
Of the mineral extraction methods, underground
mining is the most widespread in New Zealand and has
been used extensively in the coal industry, particularly on
the West Coast. Underground mines involve excavating
either an inclined shaft (a decline or drift) or a horizontal
shaft (an adit), followed by parallel shafts (bord-and-pillar
or longwalls), which enable the maximum amount of a
coal seam to be extracted. The introduction of hydromining (using high-pressure water blasted against the coal
face) improved the efficiency of coal extraction, while
exacerbating the effects of mine drainage.
Several large opencast mines have been developed,
including the Waihi (extracting gold) and Stockton (coal)
mines. Opencast mining involves the removal of surface
topsoil and rock to expose shallow mineral seams. It creates
considerable waste material or overburden, which must
Figure 36.1 Placer mine gold operation in Carton Creek
near Reefton.
Photo: Jon Harding
then be disposed of. Excavated soil and rock take up 50%
more volume than the in-place material, hence back-filling
of opencast mines still leaves substantial quantities of excess
overburden (Fig. 36.2).
Placer or alluvial mining is the other major extraction
technique used in New Zealand. Much ofthe current gold
Figure 36.2 Overburden dumps, Stockton opencast mine.
Photo: Jon Harding
36.2
Freshwaters of New Zealand
mining in the South Island is alluvial mining. Gold fines
eroded from seams in the mountains are washed naturally
down to the valley floors. These heavy gold fines filter into
the deep riverbed gravels, and alluvial mining operations
excavate the riverbed, often down to bedrock, to sift out
the gold.
Each of these three extraction methods can cause
marked changes to mining landscapes, and consequently
surface water and groundwaters associated with mines may
be significantly affected by mine leachate, sediment and
by mine operations.
In this chapter we discuss the effects of mining activities
on the water quality, physical morphology and biotic
communities of freshwater ecosystems. Increasing pressure
for improved environmental management has resulted in
a greater emphasis on the restoration and remediation of
sites affected by current and historic mining activities,
hence research is now beginning to focus on these
problems.
Much of the mining activity in New Zealand is
associated with running water ecosystems, and we will
focus on these. No major lakes in New Zealand currently
receive mine discharges.
HYDROLOGICAL EFFECTS
Many early mining activities relied heavily on a
continual supply of water, either to assist with excavation
(as in hydro-mining) or to transport mined material to the
surface. In Central Otago, for example, hundreds of
kilometres of shallow canals were constructed to transport
water from foothill streams to gold-mining digs. Initially,
this water was used for washing and sluicing gold fines.
Later hydro-mining was developed, in which high-pressure
water jets were used to carve soil and gravels from the
hillside. In addition, the manipulation of flow conditions
is common in placer mining, where flow is temporarily
diverted to expose streambeds to be re-worked for alluvial
gold. The long-term hydrological impacts of these flow
diversions have not been well studied. Prior to and during
mining the riverbed is usually cleared of vegetation and
riverbed gravels are turned over. In at least one instance on
the West Coast, channel diversion lowered the water table,
causing the diverted river to dry in late summer (Harding
and Greenwood 2003). Placer mining may also involve
diversion and storage of water in dredging and sedimentsettling ponds, which may also have significant short-term
effects on the water table.
In underground mining, the diversion of surface water
and groundwaters is also common. However, the longterm effects of altered hydrological regimes are not well
documented.
WATER QUALITY
Water chemistry associated with mining activities has
been the focus of much investigation. Surface water and
groundwaters associated with mines are often affected by
acidification, the presence of toxic metals, and
sedimentation. Discharges emanating from coal mines can
cause the most severe problems. Many coal seams, such as
those in the Brunner Coal measures in North Westland,
contain pyrite, which has high levels of sulphur. When
pyrite is exposed to water and oxygen, several well
documented reactions can occur (Singer and Strumm
1970), resulting in the acidification of mine waters.
1) FeS 2 + 7/20 2 + H 2 0 = Fe 2+ + 2S0 4 2 + + 2H +
2) Fe 2 + V4 0 2 + H + = Fe 3+ + 1/2H 2 0
3) Fe 3+ + 3H 2 0 = Fe(OH) 2 + 3H +
4) FeS 2 + 14Fe 3+ + H 2 0 = 12Fe 2+ + 2S0 4 2 + + 16H+
Thus the exposure of pyrites to oxygen is a crucial step in
the generation of acid mine runoff (sulphuric acid),
whereas in natural undisturbed coal formations, where
exposure to oxygen is uncommon, acid generation is rare.
Acid mine drainage is frequently cited as one of the most
important environmental side effects of coal mining.
In environmental monitoring of acid mine drainage,
acidity is usually assessed by measuring pH. However, pH
is not a true measure of acidity, nor is it necessarily an
accurate indicator of the extent of acid mine generation
occurring in a system (Kelly 1988). More precisely, pH is
a measure ofthe concentration or activity of hydrogen ions
in a solution. In acid generation, the crucial factor is the
availability of hydrogen ions to neutralise bases, hence, the
determining factor controlling acidity is the excess of
hydrogen ions over other ions. Thus "total acidity"
(measured as CaC0 3 ) is a more accurate measurement of
acidity than pH. However, at low acidity there is often a
correlation between total acidity and pH, and realistically
pH is easy to measure with standard field meters (Fig.
36.3). Above about pH 7 there is rarely any acidity.
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4
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7
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PH
Figure 36.3 Relationship between pH and total acidity
(after Kelly 1988).
Impacts of mining
Acidification can have severe effects on freshwater biota,
however from a chemical perspective one of the most
important side effects may be the reduction in the
bicarbonate buffering capacity ofthe water. Once pH falls
below about 4.2, carbonate and bicarbonate are converted
to carbonic acid. The upshot of this is twofold. Firstly,
stream water loses its capacity to buffer changes in pH, so
that additional acid mine effluents entering a stream may
decrease the pH markedly. Secondly, the loss of bicarbonate will have a negative impact on the many photosynthetic organisms that require inorganic carbon. Algae
and bryophytes living at pH <4.2 need to be able to process
free carbon dioxide in the absence of bicarbonate.
Furthermore, even after acid mine inputs have ceased, it
may take a stream or river many decades to recover from a
loss of buffering capacity.
Lowering of pH can also have other chemical effects.
Acidification tends to increase the rate of precipitation of
silt and clays, and thus may increase sedimentation on the
riverbed. However, depending on the pH, this may also
have the positive side effect of reducing the turbidity of
the water. Furthermore, increased acidification increases
the decomposition of minerals, including feldspars and
carbonates, resulting in the release of metals such as
aluminium, and the release of silica. The release of silica
may stimulate the growth of acidophilic algae, particularly
diatoms, whereas aluminium may have several negative
impacts. As the sources of acidity in mine drainages are
strong mineral acids, acidification in these systems is
usually associated with high conductivity (Fig. 36.4).
1000
800
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t3
600
400
36.3
Figure 36.5 — Relationship between pH, and total reactive
aluminium and dissolved iron, in 24 South Island
streams (after Winterbourn et al 2000).
acidity enhances the solubility of these metals, however as
pH increases the chemical nature ofthe metals is affected.
At very low pH, e.g., <3, any metals present remain in
solution, however, as pH increases floes of these metals
may form, and if the metals are present in high
concentrations precipitates form on the streambed. In
particular, the yellow or ochre precipitates of iron (Fe3+)
("yellowboy") are commonly found in streams where iron
is associated with exposed coal. This precipitate is
composed primarily of iron hydroxide (Fe(OH 3 )), but may
also include a complex of hydrates.
The critical pH value for iron precipitate formation
ranges from about 3.5—4.3 (Table 36.1). Below this, Fe3+
remains dissolved in stream water and although it is toxic
to some biota at high concentrations in this state, it does
not cause sedimentation problems. Iron hydroxide
precipitate can cause significant modification of the
streambed by cementing substrata, clogging interstitial
spaces, and covering biota. The substrate "armouring"
TD
C
o
O
200
Table 36.1 Minimum pH values for precipitation of metal
ions as hydroxides (Kelly 1988; Niyogi etal 1999)
Figure 36.4 Relationship between stream pH and conductivity in 24 South Island streams (after Winterbourn
etal 2000).
Mine discharges may markedly reduce the p H of
receiving waters, and in New Zealand adits with a pH <2.9
are not uncommon. A range of metals may also be
associated with coal formations, including aluminium,
arsenic, copper, iron, lead, nickel and zinc (Fig. 36.5). High
Metal
Minimum pH-hydroxide
Sn
Fe3+
Al
Pb 2 +
Cu2+
Zn
Ni
Fe2+
Cd
Mn2+
4.2
3.5-4.3
4.9-5.4
6.3
7.2
8.4
9.3
9.5
9.7
10.6
36.4
Freshwaters of New Zealand
effect caused by precipitate reduces refugia for invertebrates
and fish, making them more susceptible to floods and
disturbance and increasing the potential for predation and
competition. Other metals precipitate in streamwater at
differing pHs (e.g., aluminium), and several critical pH
values are shown in Table 36.1. When acid inputs to a
stream are stopped, these precipitates may re-dissolve and
the substrate may recover in as little as six months (Niyogi
et al. 1999). As we discuss later, Fe, Al, and other metals,
both dissolved and as precipitates, may be directly toxic to
freshwater biota.
Aluminium hydroxide precipitate, seen as a white
coating on stones, has been widely documented in North
America, whereas in New Zealand it has rarely been
reported. However, dissolved Al is a significant problem
in many catchments. On the West Coast of South Island
naturally acidic brown-water streams are common; this
acidity is produced by natural tannins (humic and fulvic
acids) derived from decomposing vegetation in temperate
rainforests. These naturally low p H waters (down to about
pH 4) may have relatively high levels of dissolved Al, which
is less toxic than in acid mine drainage waters because the
Al is bound with organic carbon (Collier et al. 1990;
Stenzel and Herrmann 1990).
Different metals have highly variable degrees of toxicity
on stream biota, depending on localised conditions. Several
workers have shown that Cu can be 13 times more toxic
than Zn; however when these two metals occur together
the cumulative toxicity can be far greater (Gray 1998).
Similarly, numerous studies have shown that dissolved Al
is more toxic than dissolved Fe at comparable
concentrations. Winterbourn et al. (2000) reported
concentrations of Fe that were three times greater than Al
in the tissues of insects from 24 West Coast Streams. Fe
can be photo-reduced by sunlight, and therefore open
streams are liable to have Fe at less toxic levels (Niyogi et
al. 1999). Not surprisingly, seasonal variations in
discharges from mines alter the concentration and toxicity
of receiving waters. Thus, a stream may be toxic at one
time ofthe year and significantly less toxic at other times.
Furthermore, heavy metals are transported in the water
column both as dissolved free metal ions and complexes,
and as metals bound to suspended sediments. In the
Waihou River, Waikato, Cu, Mn and Zn were mainly in
dissolved form in river water, whereas Fe was
predominantly bound to suspended acid-soluble
particulates in neutral pH waters (7.0-7.9). However, in
the Waihou River estuary, almost all Fe and Mn had
flocculated out, so high sediment metal concentrations
occurred there (Webster 1995).
So far we have primarily discussed conditions associated
with coal mining. However, alluvial mining may also
release minerals associated with buried riverbed material
and cause significant turbidity in receiving waters. The
successful use of a cyanide process for extracting precious
metals was crucial to the further development of gold
mining. However, the extraction process also results in
other toxic materials (e.g., arsenic, antimony) being
released. For example, X-ray diffraction analysis of
sediments showed quartz, magnetite, hypersthene,
ilmenite, anglesite, hornblende and cummingtonite were
present in sediments in the Karangahake region of the
Waihou River (Sabti et al. 2000). Waste products from
cyanide-based metal extraction include toxic sludge. In
1895, following sustained pressure from mining
companies, the Government declared (by Proclamation)
the Ohinemuri and Waihou Rivers in the Coromandel as
sludge channels (Watton 1995). This allowed the discharge
of an estimated 250,000 tonnes per annum of mine tailings
into the Ohinemuri River.
Placer gold-mining frequently results in significant
increases in suspended sediment levels downstream. For
example, turbidity has been shown to increase from 2.4 to
>100 NTUs (Nephelometric Turbidity Units) in West
Coast streams above and below placer operations (DaviesColley et al. 1992). Increased turbidity reduced light
penetration by as much as 400% and affected both benthic
algal biomass by reducing photosynthesis and benthic
invertebrate densities by degrading food quality (Quinn et
al. 1992).
MODIFICATION TO THE PHYSICAL
HABITAT
Gold dredging produces large amounts of sifted gravels,
which historically, were deposited as tailings. Numerous
West Coast and Otago rivers (e.g., the Grey, and
Taramakau) have extensive areas of modified riverbeds.
These tailings form convoluted ridge and valley hummocks
along the river banks, frequently creating small artificial
ponds, reducing or preventing riparian re-vegetation along
rivers, altering the river channel morphology, and exposing
minerals to weathering and erosion.
In contrast, opencast mining results in different, though
no less obvious, modification of the landscape.
Overburden dumps where excess pit spoil is piled can form
new hills within the mine terrain (Fig. 36.2). These
overburden dumps frequently act as sources of leachate
and sediment, which enter surface waters associated with
the mine. Capping and remediation of overburden dumps
is an emerging research challenge.
Sedimentation of waterways can be a major problem
associated with excavation and roading at mine sites, as
suspended sediment can smother algae, benthic
invertebrates and the substrate, reducing substrate
heterogeneity (Fig. 36.6).
Impacts of mining
36.5
Increased sediment
1.6
R2 = 0.48
P< 0.001
1.4
Interstitial habitats
clogged
Fish communities decline
- migration effected
- growth rates reduced
- spawning habitats impacted
Figure 36.6 Model of the pathways by which suspended
sediment might affect components of the stream
ecosystem (after Rowe and Dean 1998).
BIOTIC RESPONSES
Bacteria and fungi
Alterations in water chemistry, particularly increased
acidity and the release of metals, can have a profound effect
on
microbial
communities.
Experiments
on
decomposition rates of leaves in coal mining streams have
generally shown greatly reduced microbial activity,
although Harbrow (2001) found highly variable
breakdown rates in six West Coast streams. Winterbourn
etal (1985) observed a lack of fungi on stones in a stream
with a natural pH 4.3, while Hildrew et al (1984) noted
impaired cellulolytic decomposition in streams with pH
<5.6. The mechanisms that lead to a reduction in microbes
at low pH are unclear, but there are at least two. Firstly,
high acidity inhibits microbial enzymes. For example, the
activity of pectin-degrading enzymes such as pectin lyase,
which is involved in softening plant tissue in the early
stages of leaf breakdown, is lower at low pHs (Suberkropp
and Klug 1981). Secondly, reduced litter breakdown rates
have been linked to metal oxide deposition, and Niyogi et
al (2001) found that microbial respiration decreased with
increasing rates of deposition of metal oxides on leaf litter
(Fig. 36.7). Thus, the precipitates themselves may either
smother microbial complexes or be directly toxic to them.
Consequently, in lakes and rivers affected by mining
and not subject to flushing flows from floods, coarse
particulate organic matter may accumulate on the bed
because natural decomposition is suppressed. Reduction
in microbial processing of allochthonous inputs may also
significantly reduce the availability of dissolved and fine
organic matter within the system. These interrelationships,
when coupled with a reduction in the numbers of
invertebrate shredders due to toxic conditions, can
combine to reduce overall allochthonous breakdown in
mine drainage streams (Fig. 36.8).
O.O
0.01
0.1
1
10
Deposition rate of metal oxides (g m" d")
Figure 36.7 Microbial respiration and deposition rate of
metal oxides (after Niyogi et al. 2001).
The importance of metal oxide deposition has been
confirmed in post-remediation studies in which litter
breakdown rates failed to recover even after water quality
improved. In some instances the presence of metal oxides
may have continued to limit microbial activity. Poor or
depauperate food resources frequently have been cited as
additional confounding factors in the recovery of benthic
invertebrate communities in mine drainage streams after
remediation, however this has not been well studied in
New Zealand.
The presence of high concentrations of dissolved metals
has little effect on some species, particularly iron bacteria
that occur in some systems affected by mining.
Winterbourn et al (1985) found the iron bacterium
Leptothrix sp. dominated a stream with total iron
concentrations of 6 g rn-3, and was eaten by a chironomid
that lived on it. Little is known about the ability of other
organisms to use these food resources.
Dissolved
metals
Mine
drainage
Low pH
Metal oxide
deposition
Shredder
biomass
. Microbial
activity
^^w
^ k
Litter
breakdown
FPOM
/ *
Figure 36.8 Model ofthe effects of mine drainage on leaf
litter breakdown and the production of fine particulate
organic matter (FPOM). The thickness ofthe arrow
indicates the strength ofthe effect (modified from Niyogi
etal 2001).
36.6
Freshwaters of New Zealand
Aquatic plants
Vascular plants are usually absent from acidic streams,
but bryophytes and periphyton can be locally very
abundant. Some algae are acidophilic and can occur in
high densities in acidic sites with stable flows. Where acidic
algae proliferate, communities are frequently dominated
by a few species, with high biomass. Acid-tolerant algae,
such as Ulothrixsp., have been shown to account for 99%
of the algal biomass at mine discharges, with other
filamentous algae such as Microspora, and Tribonema
common (Niyogi et al. 1999; Winterbourn et al 2000)
(Fig. 36.9).
1996). It has been suggested, however, that Fe can actually
decrease the toxic effect of other metals by competing for
binding sites, and by the partial sequestration of other
metals by Fe colloids. Al has also been shown to accumulate
in bryophytes, however Al absorption in moss usually
peaks in tissue at intermediate pHs, i.e., 5.2-5.8
(Englemann and McDiffett 1996; Winterbourn et al
2000). The reason for this is probably that the solubility
ofAl increases markedly below p H 5, where the more toxic
Al3+ is more prevalent. Hence the higher concentration of
Al in plants at intermediate p H may be a result of
bioaccumulation during periods of lower pH, and poor
solubility as p H rises. The low accumulation of Al at low
pH may be due to the domination of Al3+, which has been
shown to cause mucilage production and root necrosis in
duckweed (Crowder 1991). Where they are able to survive,
higher-order aquatic plants can also accumulate metals.
Sabti et al (2000) found that aquatic macrophytes {Egeria
densa) in the Ohinemuri and Waitekauri Rivers contained
gold (302 and 672 jug kg"1 dried matter respectively), and
suggested that plants may be useful as bioindicators of
heavy metal contamination.
Benthic invertebrate communities
Figure 36.9 Algae and iron hydroxide precipitates at
Sullivans West adit, Denniston Plateau, North Westland
Photo: Jon Harding
Like microbial communities, algae may proliferate at
sites with low pH, stable flows and low metal oxide
deposition (e.g., mine adits), but as soon as metal
deposition rates increase (e.g., when pH rises above 3.5—
4.3) algae can survive only if they are able to grow faster
than the rate at which oxides smother their surfaces.
Experimental manipulations of pH have shown that algal
biomass can increase until Al precipitation occurs (at about
pH 4.9) and periphyton growth is inhibited.
Where the substratum is relatively stable, bryophytes
(mosses and liverworts) occur in naturally and miningaffected low-pH streams. The liverworts Lophocolea,
Jungermannia and Riccardia, and mosses such as Blinda
and Sphagnum, have been recorded in low p H streams in
NewZealand (Winterbourn etal 2000). Some bryophytes
accumulate metals and, in particular, Fe. Winterbourn et
al (2000) reported concentrations of Fe 10 times higher
than Al in bryophyte and algal tissue in West Coast
streams. Concentrations of metals can reach high levels;
Englemann and McDiffett (1996) reported concentrations
of 17.3 and 9.1 jlg/g dry mass of Fe in two species of
bryophytes. Fe3+ is practically insoluble, and moderately
toxic to plants. It competes with other substances for
binding sites on cell membranes and is then taken into the
cell where it accumulates (Englemann and McDiffett
The impacts of mine drainage on stream invertebrates
are almost entirely negative and have been termed
"acidaemia" by Kelly (1988). The effects range from acute
and direct toxicity caused by combinations of low pH and/
or the presence of toxic metals, to indirect effects in areas
where food resources are limited (by reducing organic
matter processing and algal growth), to altered in-stream
habitat (by armouring and clogging of the riverbed
substrata).
In naturally acidic brown-water streams (pH 4.3-5.7)
benthic invertebrate communities often have fewer species,
lower densities and altered community composition
compared to communities in similar-sized natural streams
(Collier and Winterbourn 1987). Collier and Winterbourn (1987) suggested that the depauperate state of
naturally acidic brown-water streams was probably a result
of changes in the food supply, particularly a reduction in
periphyton biomass (dominated by diatoms such as
Eunotia and Fragilaria spp.). By comparison, streams
affected by acid mine drainage may be almost devoid of
species (Fig. 36.10), and where organisms are present their
densities may be as low as a few animals per square metre.
Naturally acidic waters are not uncommon on the West
Coast of New Zealand, and it is apparent that some benthic
invertebrate species are well adapted to the conditions
found in low p H waters (Winterbourn and McDiffett
1996). Naturally acid brown-water streams in South
Westland typically have communities dominated by
the common leptophlebiid mayfly
Deleatidium,
Chironomidae, the elmid beetle Hydora, and the stonefly
Impacts of mining
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Figure 36.10 Relationship between total taxa and pH in
23 streams in South Island (after Winterbourn et al.
2000).
Zelandobius. However, in streams contaminated by acid
mine drainage in North Westland (pH 2.6-4.2), a range
of taxa were recorded in low densities, including the
mayflies Deleatidium and Austroclima sepia, the stonefly
Spaniocercoides philpotti, the caddisflies Psilochorema and
Oxyethira albiceps, scirtid beetles and dipterans, including
Chironomus zealandicus and Eukiefferiella (Boothroyd
2002; Harding 2002). Sites with pH <3 in the Denniston
and Stockton Plateaus were dominated by chironomids
and scirtids, but included two caddisflies—Pseudoeconesus
and Kokiria miharo (Winterbourn 1998). Although
Crustacea are frequently considered intolerant of acidity,
Anthony (1999) found the amphipod Paraleptamphopus
below p H 4, and the crayfish Paranephrops occur in
naturally low pH waters (Collier et al 1990).
Exposure to low pH and elevated metal concentrations
may cause a number of physiological stresses to benthic
invertebrates. Numerous studies have shown osmoregulation in many species is impaired by the disruption
of acid base and ion balances, nitrogen excretion and
respiration (Sutcliffe and Hildrew 1989). Low p H also
causes a reduction in sodium uptake, detrimental to
sodium-sensitive species such as crayfish (Haines 1981).
Similarly, Cl, Ca and K can all be lowered in high-acidity
waters (Rowe etal 1989), while Bell (1971) showed that
low pH could be lethal at critical junctures ofthe life cycles
of stream insects. Thus, trichoptera adults had only 50%
successful emergence at pH 4, while mayflies showed 50%
adult emergence failure at pH 5.9. However, a number of
New Zealand mayflies and caddis are clearly capable of
successful emergence at low pH, although this may differ
between natural and mine waters. The dominance of
benthic communities by Chironomidae at low p H may
relate to the ability of these taxa to tolerate ion imbalance
(Forsyth 1983; Boothroyd 2002).
36.7
Low pH may also affect some species by making them
more susceptible to disease, infection and parasitism.
Leuven et al. (1986) found that amphibian eggs were
unable to develop at p H <3.5 due to infection from
pathogenic fungi, while the Crayfish Orconectes virilis
suffered egg mortality from protozoan parasites when the
pH dropped below 5.6 (Schindler and Turner 1982). A
high frequency of infections (visible sores) has been noted
in benthic invertebrates from streams receiving discharge
from abandoned gold mines near Reefton. This suggests
that toxins in these mine waters may weaken benthic
organisms, making them susceptible to disease and
infection.
Toxicity of metals can be a major problem in many
mining-affected waters. Toxocological studies have shown
insects, Crustacea and fish are frequently susceptible to
toxic metals. Hickey and Clements (1998) noted that netspinning hydropsychid caddis and orthoclad chironomids
dominated at sites high in metals in the Coromandel
Peninsula. Analysis of taxonomic richness data from North
Westland shows that relatively few species are found in
waters with high concentrations of dissolved Al and Fe
(Fig. 36.11).
Al (mg I"1
40
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3 0 -j
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Fe(mgr )
Figure 36.11 Relationship of total taxa to mean total
reactive aluminium (top) and total dissolved iron
(bottom) in 19 streams in North Westland (after
Harbrow2001).
36.8
Freshwaters of New Zealand
Fish communities
Table 36.2 Mean copper concentrations in muscle tissue of rainbow trout
from the Waitekauri River, Golden Cross (All data stated as mg/kg wet
weight).
Most fish species are negatively affected
by acidification and mine leachate. These
toxic effects may be acute—causing death,
Immediateiy
5 km
Control
or chronic—resulting in impaired health
Upstream
Year
downstream downstream
site
(e.g., mucous secretion on gills impairing
1993
1.22 + 0.29
1.39 ±0.09
0.62 ± 0.08
gas exchange), or physiological—from
1994
1.15±0.14
0.76 ± 0.07
0.64 ± 0.04
stress that reduces fish condition.
1996
0.32 ± 0.08
0.31 ± 0.06
0.41 ±0.15
0.29 ± 0.04
Accumulation of heavy metals within the
flesh of fish may have long-term toxic
NTUs (Boubee et al 1997). A possible explanation for
effects (Table 36.2). Furthermore, the presence of mine
these preferences was offered by Rowe and Dean (1998),
discharges may create a chemical barrier to diadromous
who
demonstrated that feeding rates of banded kokopu
species, reducing or preventing their migration to and from
and inanga declined significantly as turbidity increases.
the sea. Several diadromous native fish have been recorded
Banded kokopu feeding rates were 40% lower at 20 NTUs
in low pH waters (Main 1988) and have been shown to be
than in clear water, whereas koaro feeding rates were not
able to detect pH gradients and display p H preferences. In
affected, even at the highest turbidity.
particular, short-finned eels {Anguilla australis) and two
whitebait species, koaro (Galaxias brevipinnis) and banded
kokopu {Galaxias fasciatus), prefer waters <6.5, whereas
Ecosystem interactions
other species avoided low pHs in laboratory trials (West et
As we have seen, mining activities can have a profound
al 1997). These pH preferences were stronger in adults
effect on water chemistry, physical conditions and the biota
than juveniles, and probably account for the presence of
of freshwater systems. The loss or reduction of bacteria
galaxiids in naturally brown-water streams on the West
and shredding insects means that the processing of organic
Coast. The freshwater shrimp Paratya curvirostris seemed
matter is significantly reduced in mining-affected streams.
unable to detect pH changes, and had high mortality in
Similarly, impacts on algal communities, which form the
high pH waters. These findings are consistent with Collier
food base of many New Zealand stream food webs
et al (1990), who reported widespread tolerance of low
(especially outside forests), will have a cascade effect on
p H in naturally acidic streams in Westland, with 9 out of
invertebrate grazers, predators and fish communities.
14 native fish species reported in p H <5, and 7 species at
Figure 36.12 summarises the major chemical, physical and
p H <4.5. Particularly low pH-tolerant species were inanga
biological responses from the effects of mining.
{Galaxias maculatus), giant kokopu {G argentus), long- and
short-finned eels and banded kokopu. The fish fauna in
mine-contaminated streams have not been
well studied. Historical anecdotal evidence
Reduced pH
from local residents in the Ngakawau River,
Increased
turbidity
north Westland, indicate that prior to
mining crayfish and koaro were abundant
Metal adsorption
in this system, however neither species have
on substrate
Increased soluble
been recorded post-mining.
metals
As mentioned previously, the presence of
high concentrations of suspended
sediments derived from mining (e.g., from
placer mining, roading) can affect fish
communities. Behavioural experiments by
Boubee et al (1997) on native juvenile
migratory species have shown that banded
Habitat
Elimination
loss
kokopu can be sensitive to turbidities >17
of species
NTUs, whereas koaro and inanga are less
Reduction in food
Bioaccumulation of metals
sensitive and avoid only much more turbid
quality and quantity
silt-laden waters (>70 and 240 NTUs,
respectively). In contrast, short-finned and
long-finned eels, and red-finned bullies Figure 36.12 Model ofthe primary ecosystem factors influenced by mine
drainage inputs (modified from Gray 1997)
were not effected by turbidities > IOOO
Impacts of mining
Management and restoration strategies
DAVIES-COLLEY, RJ.; HICKEY, C W ; QUINN, J.M.; RYAN,
The chemical and geological complexity of the effects
of mine drainage presents significant challenges to the
effective management of mining landscapes. The nature
ofthe industry (i.e., large-scale extraction of minerals from
the ground) means that environmental impacts are a
byproduct of the process, and moderating these impacts
has become the focus of research. Management strategies
centre on reducing toxic inputs, remediation of effects,
restoration and monitoring. While a number of
remediation techniques have been developed internationally, the success of many techniques—artificial
wetlands, limestone dosing, dilution and buffering, and
bacterial reduction of leachate—have been highly variable,
and have not been adequately tested in New Zealand.
SUMMARY
Considerable knowledge exists about the chemical
processes involved in mine drainage, and its effects on
water chemistry, algae and benthic communities have been
well documented. The complexity of these relationships
continues to present significant challenges to researchers,
and those charged with the task of remediating mine
impacts.
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